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The progressive downward shift in dominance of key reef building corals, coupled with dramatic increases in macroalgae and other nuisance species, fields of unstable coral rubble ,loss of structural relief, and declines of major functional groups of fishes is a common occurrence throughout the Caribbean today. The incorporation of resilience principles into management is a proposed strategy to reverse this trend and ensure proper functioning of coral reefs under predicted scenarios of climate change, yet ecosystem processes and functions that underlie reef resilience are not ]]>
M. ]]>
(complex) skeletons. Nevertheless, completely dead M. annularis colonies were common, survivors were frequently reduced in size and subdivided into smaller tissue remnants, and these species exhibited higher amounts of partial mortality than all other species. A notable absence of sexual recruits and juveniles of M. annularis illustrates a progressive shift away from a Montastraea dominated system. This shift, characterized by an ]]>
Agaricia and Porites and a reduction in size of longer-lived massive corals, is occurring throughout the Caribbean. Monitoring of the survival of recruits is necessary to determine whether Caribbean reefs will retain the same function, structure, identity and feedbacks (key signs of resilience) if the losses of M. annularis (complex) continue at present levels. The rapid assessment protocol utilized here allows ]]>
Key words: resilience, coral size structure, coral recruitment, survival and growth, coral monitoring and assessment.
Resumen
En la actualidad se está viendo en el Caribe un cambio en la composición de los corales constructores de arrecifes, aumento en la cobertura de macroalgas y otras especies, un aumento en áreas cubiertas por escombros de ]]>
Montastraea annularis (complejo). Colonias completamente muertas de M. annularis fueron comunes y los sobrevivientes con frecuencia son más pequeños o subdivididos en pequeños restos de tejido. Montastraea annularis es la especie que exhibe una mayor mortalidad ]]>
M. annularis ilustra el cambio progresivo de cambio de un sistema dominado por Montastraea. Este cambio, que se está produciendo en todo el Caribe, se caracteriza por un dominio cada vez mayor de especies más pequeñas y de vida corta como Agaricia y Porites, y una reducción en el tamaño de los corales masivos longevos. ]]>
M. annularis (complejo) continuarán a los niveles actuales. La evaluación rápida presentada aquí posibilita caracterizar la estructura de tamaño de las colonias, los niveles de reclutamiento y determinar si los corales pequeños ]]>
Palabras clave: resiliencia, estructura en la talla del coral, reclutamiento de coral, sobrevivencia y crecimiento, monitoreo ]]>
Until the late 1970s, benthic substrates on Caribbean reefs were occupied primarily by reef-building corals, with 50-80% benthic cover by living corals. Coral reefs exhibited a generalized zonation pattern with elkhorn coral (Acropora palmata) forming large, monospecific stands in the reef crest ]]>
A. cervicornis) at intermediate depths (5-25 m depth) on wave exposed reefs and in shallow, protected environments; massive corals (dominated by Montastraea annularis complex) throughout the fore reef (5-30 m depth) and in back reef and lagoonal areas; and plating agaricids near the base of the reef (20-40 m depth) (Goreau 1959). Coral cover has ]]>
et al. 2002, Bruckner 2003, Gardner et al. 2003), which is now being followed by massive losses of Montastraea annularis (complex) (Bruckner & Bruckner 2003, 2006a,b, Edmunds & Elahi 2007, Bruckner & Hill 2009). ]]>
Changes to Caribbean reefs, attributed largely to coral diseases, hurricanes, mass mortality of the herbivorous long-spined sea urchin (Diadema antillarum), localized human impacts, recent bleaching events and climate change (Lessio 1988, Carpenter 1990, Sutherland 2004, Weil 2004,Grimsdith & Salm 2006) have manifested as dramatic phase shifts characterized by a dominance of macroalgae and other nuisance species, fields of unstable coral ]]>
Agaricia and Porites (Hughes 1994, Edmunds & Carpenter 2001). Many of these stressors have cumulative negative impacts on coral reef ecosystem health and are integrally linked. For example, the virulence and severity of diseases is elevated during and immediately following bleaching events and other periods of elevated environmental stress (Bruno et al. 2007, Ballantine ]]>
et al. 2008, Muller et al. 2008, Rogers et al. 2008, Bruckner & Hill 2009). Other human impacts such as overfishing of groupers, snappers, parrotfish, lobster and other species can have cascading impacts on ecosystem health by removing keystone species that are critical in controlling harmful algae, corallivores and other pests (Jackson et al. 2001, Mumby 2006, Green & Bellwood 2009). These “pest” species compromise remaining corals through direct competition ]]>
Although many of the threats affecting coral reefs have been examined in detail since the early 1990s, very few studies have identified a single direct cause of these impacts (a “smoking gun”), or site-specific remedies or treatments for particular problems. Conversely, there has been a proliferation of ]]>
et al. 1996, Roberts 1997, Roberts et al. 2001, Hughes et al. 2003, 2005, Almany et al 2007, 2009, Jones et al. 2009,.. There is good evidence that reduced fishing pressure, and in particular the protection of herbivorous fish populations, can prevent ]]>
et al. 2006). Furthermore, restoration of certain keystone invertebrates, such as Diadema antillarum, and increases in density of large herbivorous fishes can trigger a reversal from a macroalgal dominated state and promote coral recruitment (Edmunds & Carpenter 2001, Carpenter & Edmunds 2006, Hughes et al. 2009, Norstrom et al. 2009). As ]]>
Over the last decade, research efforts have placed a greater emphasis on examining coral reef health and resilience, with a goal to identify strategies to enhance the resilience of these ecosystems (Hughes
et al. 2003, 2005). Ecological Resilience is a term which describes the capacity of a system to absorb, resist or recover from disturbance or to adapt to change, while continuing to maintain essential functions, structure, identity and feedbacks (Hollings 1973, Nystrom & Folke 2001, Nystrom et al. 2008, Obura & Grimsditch 2009). There is considerable evidence that sites with chronic human impacts are the least likely to resist mortality and subsequently recover from acute large-scale events, but ]]>
et al. 2006, Knowlton & Jackson 2008). MPAs, when properly designed and enforced may help promote resilience to disturbances by increasing or maintaining key ecosystem parameters such as fish biomass and coral cover, and help maintain critical ecosystem processes and functions (Cote et al. 2001, Halpern 2003). However, in some cases MPAs may not ]]>
et al. 2006), emphasizing the need for alternate management strategies. While great strides have been made in understanding interrelationships among corals, fishes, and algae and the role of certain functional groups in enhancing the health of coral reef ecosystems, large gaps remain in our understanding of the dynamic and complex processes that promote or undermine resilience (Hughes et al. 2005). ]]>
A detailed assessment of resilience relates to the entire scope of positive and negative factors affecting the ecosystem. This includes ecological, environmental and physical factors as well as social factors such as patterns of resource uses and extraction, type and extent of pollution, presence of invasive species, governance structures, economics, and effectiveness of existing conservation and management efforts (Hughes et al. 2005). Quantitative ]]>
et al. 2006, Lindsey & Bruno 2008, Nystrom et al. ]]>
The Atlantic and Gulf Rapid Reef Assessment (AGRRA) and the IUCN Resilience Assessment protocols (Obura & Grimsdith 2009, Lang et al. 2010) are two different rapid ecological assessment approaches that can provide an indication of the ]]>
ecological resilience of a coral reef. Both of these share common attributes, and they highlight the importance of the inclusion of ecologically relevant fishes, algal functional group biomass/cover, and coral population structure in rapid assessments. The IUCN protocol includes other measures as well, such as a qualitative assessment of various ecological and biological factors as well as physical factors that promote resilience through shading, screening, cooling and enhanced stress tolerance (Obura & Grimsditch 2009). A detailed resilience assessment also includes characterization of reef processes, ]]>
Because comprehensive measurements of resilience, incorporating the parameters described above and others, ]]>
Materials and Methods
Study site: In July 2011, the Living Ocean ]]>
Montastraea dominated reefs from 5-15 m depth. All of these sites are fringing reefs that begin close shore and drop to deeper water within a few hundred meters. Many of the areas above 5 m depth were badly damaged by recent hurricanes and are largely devoid of corals, but much of the Montastraea structure at mid depths is still intact. The sites were grouped into 3 distinct areas: a) sites north of Kralendijk located off ]]>
Fig. 1, Table 1).
Assessment protocol: The resilience ]]>
Benthic cover: Cover of substrate and benthic organisms (algae and invertebrates) was estimated using a point intercept method. At each site, a minimum of six 10 meter long transects were deployed. The organism and substrate type was recorded every ten cm for a total ]]>
Microdictyon, Lobophora, Dictyota, Stypopodium, Peyssonnelia). Key invertebrates that also may become nuisance species recorded to genus ]]>
Trididemnum), encrusting gorgonian (Erythropodium, Briareum), colonial anemone (Palythoa), encrusting or bioeroding sponge (Cliona langae/aprica complex, Cliona delitrix, Anthosigmella), and hydrozoan coral (Millepora).
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Recruitment: Sampling for corals smaller than 4 cm was done using a minimum of five 0.25m2 quadrats per transect. Each quadrat located at fixed, predetermined intervals (e.g. 2, 4, 6, 8, 10 m), alternating between right and left side of the transect. Recruits were identified in both point intercept surveys and belt transects. These corals were divided into recruits (0-2 cm diameter) and juveniles (2.1-3.9 cm).
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Corals colonizing skeletons of other corals: All corals settling on skeletal surfaces of other colonies (completely dead corals and corals with partial mortality) within each belt transect were identified and recorded separately from those corals occurring on reef substrates. For each of these colonies, measurements were taken of the size (length, width and height for colonies 4 cm or larger; maximum diameter for corals that were 0-3.9 cm), and an estimate of percent mortality made for those exhibiting partial mortality.
Condition of corals: Visual estimates of tissue loss, using a 1 m bar marked in 1 cm increments, was recorded for each colony over 4 cm in diameter. If the coral exhibits recent tissue loss, the amount of remaining tissue, the percent that recently died and the percent that died long ago were estimated for the entire colony surface. Tissue loss was categorized as recent mortality ]]>
Hermodice carunculata). Lesions were then diagnosed into four ]]>
et al. 2008, and included yellow band disease (YBD), white plague (WP), black band disease (BBD), red band disease (RBD), Caribbean ciliate infection (CCI), dark spots disease (DSD) and white band disease (WBD). ]]>
Results
Coral cover: On most of the reefs examined in this study, between 5-15 m depth, coral cover was high ( 40-60%;Fig. 2a). The only exceptions were two sites on northern reefs, ]]>
Montastraea annularis (complex) were the dominant corals, in terms of living cover, occupying approximately 20-25% of the benthos, and making up over 50% of the total live coral cover. The next most common species, in terms of living cover, where Agaricia, Madracis and Porites spp (Fig. 2b). Cover of major functional groups of reef building corals showed slight ]]>
M. annularis complex was lowest on northern reefs (pooled for all depths; 22% vs 24%) while cover of Agaricia spp. was lowest on southern reefs (6.6% vs 9%). Klein Bonaire also had a higher cover by Madracis mirabilis (>6%); this coral formed large thickets on several reefs ]]>
Agaricia spp. and Eusmilia fastigiata had the highest cover on all reefs at 15 m. In contrast,cover of Porites was higher at 10 m and 15 m depth than at 5 m. Madracis spp. Was most variable, having the highest cover at 5 m depth on northern and southern reefs, and 10 m depth on ]]>
Acropora palmate and A. cervicornis were virtually absent from all point intercept surveys, and only identified infrequently in belt transects. Cover of macroalgae was relatively low on all reefs, but was significantly higher on northern reefs (Fig. 2a). There was no correlation between coral cover and macroalgal cover (r2=0.01, p=0.56). Cover of turf algae ranged from 7-38%, and was significantly correlated to coral cover (r2 = 0.44, ]]>
Coral composition: A total of 5957 corals, 4 cm or larger in diameter, were identified within belt transects (10 m X 1 m) on 25 reefs in Bonaire (Fig. 3). M. annularis (complex) was the dominant functional group of corals at all sites overall, in terms of numbers of colonies, making up approximately 27% ]]>
M. annularis complex was numerically dominant between 5-10 m depth, while Agaricia was slightly more abundant at 15 m depth. Reefs in the north and off Klein Bonaire had a higher proportion of M. annularis (complex) colonies (30 and 29% respectively), when compared to reefs in the south (21%). When examined by species, M. annularis was ]]>
M. faveolata and M. franksi on northern reefs and Klein Bonaire, but not southern reefs, while abundance of M. faveolata and M. franksi was very similar in all locations. The second most abundant functional group was the genus Agaricia (18-26% of all corals). This genus was the dominant taxon on ]]>
Porites was the third most abundant taxon. While the proportion (number of colonies) of brooding species (especially Agaricia, Porites) was very high, the contribution to living coral cover was less than M. annularis (complex) because these colonies were smaller in size. ]]>
Coral size structure: Corals ranged in size from 4 cm (smallest coral assessed in belt transects; 0-3 cm corals assessed separately using quadrats and on exposed skeletal surfaces) to over 450 cm diameter, with a maximum height of 460 cm. The size structure of M. annularis ]]>
Fig. 4a). Colonies of other species (all species except M. annularis complex) were dominated by very small colonies (<20 cm) and populations in all locations (all ]]>
M. annularis (complex) were significantly larger than all other species (mean diameter = 58 cm, versus 24 cm for other species pooled; original diameter of the colony; Fig. 4b), although large (> 1 m diameter) colonies of Siderastrea siderea, Stephanocoenia intersepta, ]]>
natans and extensive (2-5 m wide) thickets of Poritesporites and Madracis mirabilis were seen.
There was a notable absence of smaller colonies (<10 cm) of M. annularis (complex). It is important ]]>
Montastraea that had contiguous skeletons,1-3 m in diameter or larger, were reduced in live tissue area, and individual corals often consisted of multiple tissue remnants (mean = 6.6 remnants per colony) that were reduced to a few cm in diameter.
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Colony mortality: The amount of partial mortality observed on corals located within belt transects varied from 0-99%, with significant differences between species, colony sizes and locations (Fig. 5). For M. annularis complex (n=1602), a total of 73 (4.5%) had completely died, with surviving colonies (n=1529) missing a mean of 28% of their tissue. Tissue loss for these species consisted of ]]>
M. annularis (complex) with less than 30% partial tissue loss were significantly smaller in size (mean diameter = 48 cm; mean tissue loss=11%; n=889) than colonies with 30-99% partial tissue loss (mean diameter =61 cm; mean tissue loss= 50%; n=639). However, a correlation analysis using the entire range of sizes (no pooling into size classes) showed that tissue loss was not significantly correlated with colony size. This may be due to the fact that colonies exhibited a high range of tissue loss in all size ]]>
M. annularis (complex) were also frequently divided into a number of smaller patches of live tissue; on average, each coral was subdivided into 6.6 separate tissue remnants.
Tissue loss for ]]>
M. annularis complex) was significantly lower (mean partial tissue loss=8%) and fewer dead colonies were identified (n=20, 0.4%). Interestingly, partial mortality for colonies that had colonized reef substrates was higher than partial mortality for colonies that had colonized exposed skeletal surfaces of living corals (9.4% vs 7.6%). These corals were also less frequently subdivided into smaller tissue remnants (mean=1.41 remnants/colony), possibly because they were younger and smaller in size overall.
Based on size structure, abundance, levels of recruitment, and coral condition, coral communities could be divided into two primary groups, the M. annularis complex (M. annularis, M. faveolata and M. franksi) and all other species. Corals ]]>
]]>
Recruits on reef substrate: A total of 1688 quadrats, each 0.25 m2, were examined on 25 reefs in Bonaire. Over 40% of the quadrats contained at least one coral (0-3 cm in diameter, classified here as a recruit). A higher proportion of quadrats on southern reefs contained recruits (45% vs. 40% and 37% of quadrats on northern reefs and Klein Bonaire, respectively). Individual quadrats contained a maximum of 5 recruits (northern and southern reefs) and 8 ]]>
M. annularis, M. faveolata and M. franksi, although numerous tissue remnants <4 cm in diameter were noted. The dominant corals observed as recruits included A. agaricites (mean = 3.9 recruits/m2), P. astreoides(mean = 2.2 recruits/m2), and Madracis spp (mean = 0.95 recruits/m2) while all other species were at densities of <0.2/
m2 (Fig. 6a). Species that were common as adults within transects, but not observed in quadrats included Mussa angulosa, Mycetophyllia lamarckiana, M. aliciae, Dendrogyra cylindricus, Isophyllia sinuosa, I. rigida, S. intersepta, Scolymia spp., Acropora palmata, and A. cervicornis. ]]>
Colonization of coral skeletal surfaces: The exposed skeletal surfaces of 249 corals were colonized by new stony corals. Corals that supported these settlers were missing a mean of 60% of their tissue. Difference in partial mortality occurred between regions, with colonies from southern reefs that were colonized by other corals showing much higher amount of partial mortality overall.
M. annularis (complex) colonies that supported settlers of other species were similar in size in all three locations, and they were significantly larger than all other species that were colonized by new corals (Table 2).
A total of 12 species of corals ]]>
(Fig. 7), although most were observed on M. annularis (36%) and M. faveolata (35%). A much higher proportion of M. annularis (complex) skeletons on Klein Bonaire (19%) were colonized by corals as compared to northern (13%) and southern (5.3%) reefs. Exposed skeletal surfaces of other species were less frequently colonized by ]]>
M. cavernosa and S. siderea on southern reefs (Fig. 4). No differences in the number of settlers per unit area were noted between reef substratum and exposed M. annularis skeletal patches.
Overall, 825 corals ]]>
Agaricia agaricites (43%) and P. astreoides (27%) were the most common colonizers, although broadcast spawners were also observed (Fig. 6b). Interestingly, some species observed as recruits were not ]]>
M. decactis and higher percentage of colonizers of E. fastigiata, C. natans, Diploria spp., M. cavernosa, P. Porites, M. meandrites, and S. interepta), but not all species (Agaricia and
P. astreoides). There was also a notable absence of M. annularis (complex) recruits and colonizers on dead skeletons. While M. annularis (complex) was not observed recruiting onto any exposed skeletal surfaces, juvenile corals (4 cm or larger) that had settled on dead coral surfaces represented 17.5% of all corals of other species (all species except M. annularis) identified within belt transects.
Discussion
Assessment protocols for stony corals frequently rely on single or repeat assessments of coral cover as the primary metric to characterize reef condition and gauge changes. ]]>
This approach assumes that reefs with high coral cover are in better shape and exhibit higher resilience (Gardner et al. 2003, Bruno & Selig 2007), yet coral cover may actually be one of the last metrics to indicate ecosystem failure (Bellwood et al. 2004, McClanahan et al. 2011). A second coral metric ]]>
et al. 2008). These parameters alone may fail to provide a reliable measure of reef resilience because they fail to take into account factors affecting the fitness and dynamics of coral ]]>
Two of the most important parameter to consider when assessing resilience of coral populations is size and amount of partial mortality, as mortality and fecundity rates in corals are strongly size-dependent (Harrison & Wallace 1990). Smaller ]]>
et al. 1994).
Since colony size can influence patterns of mortality and fecundity, population size frequency distributions provide insight into the effects of disturbance and population processes. Coral populations are normally highly positively skewed, with a dominance by the smallest size classes and an exponential decrease with increasing colony size (Bak & Meesters ]]>
et al. 1994). One consequence of this is that smaller colonies suffer total mortality more frequently than larger colonies, while larger colonies sustain higher levels of partial mortality over multiple disturbances, but have a larger chance of survival (Soong ]]>
Scleractinian coral populations normally exhibit a positively-skewed (left) size frequency distribution with populations dominated by smaller colonies. On a Caribbean ]]>
Montastraea reef, a population that has not been exposed to a major disturbance for several decades may progressively exhibit a shift towards the opposite extreme, with a dominance of extremely large colonies (negatively-skewed right). In Bonaire, a northern reef (Taylor Made) with a higher coral cover than all other sites examined in this study was dominated by extremely large colonies of Montastraea faveolata and M. annularis, many over 200 cm in diameter and 5 m in ]]>
The demographic structure of a coral population also illustrates the presence of past disturbances, where a population dominated by small size classes may indicate high mortality in intermediate and larger corals (Nystrom et al. 2008). Several reefs at the southern end of Bonaire had high coral cover, but ]]>
Agaricia and Porites, although various brain corals, flower coral, cactus corals and many other species were present. Most Montastraea colonies were small (<90 cm diameter), many dead standing colonies of these species were observed, and colonies were frequently subdivided into small tissue remnants. Although high numbers ]]>
Montastraea colonies below 10 cm diameter were observed. The declining condition of M. annularis is indicative of a past disturbance, as well as chronic tissue loss from disease. While coral cover is rebounding, the progressive replacement of M. annularis may be indicative of lowered resilience as the community is becoming dominated by smaller, ]]>
By subdividing mortality into three categories, recent, transitional and old mortality, the cause of tissue loss and the timing of a disturbance event may be discernable(Lang 2003). The presence of high levels of recent mortality suggests that a disturbance has occurred and the event is ongoing. Sites containing corals with extensive transitional mortality and little recent ]]>
The challenge faced when assessing colony size and extent of mortality through a single rapid assessment is that it is impossible to determine the growth of a coral population over time. Because corals are clonal animals, they can progressively increase in size with age, theoretically with indeterminate growth, but ]]>
In Bonaire, corals ]]>
Montastraea annularis (complex). The majority of these corals were completely alive, unlike many of the similar sized corals that occurred on the reef substrate. There were however, no differences noted in the number of recruits recorded on coral skeletons of M. annularis, versus the ]]>
Montastraea may be optimal, due to high rates of survival and growth into larger size classes (and absence of partial mortality as observed among the same species on reef substratum). Because the settlement surface (coral skeleton) is raised off the bottom, settlers are less affected by sediment transport and siltation, macroalgal competition, and possibly disease. Nevertheless, one species (Siderastrea siderea) ]]>
The discrimination of corals settling onto skeletal surfaces of other corals provides a useful metric when using a single rapid assessment to determine levels of ]]>
M. annularis (complex) exhibiting slow, chronic mortality from yellow band disease (e.g. areas that were denuded 30-90 ]]>
The results ]]>
M. annularis complex) on these ]]>
Montastraea-dominated system to a community consisting predominantly of other species. Fortunately for Bonaire, these reefs still have unusually high coral cover, high levels of recruitment and good survival of juvenile corals, and several sites ]]>
Montastraea annularis (complex), all which are important indicators of resilience.
While single assessments will never provide the kind of data that would result from repeated visits to a site over time, single rapid assessments can provide valuable data on the resilience of coral populations. For corals, these assessments must ]]>
Acknowledgements
The intensive surveys completed In ]]>
References
Almany, G.R., M.L. Berumen, S.R. ]]>
coral reef fish populations in a marine reserve. Science 316: 742-744. [ Links ]
Almany, G.R., S.R. Connolly, D.D. Heath, J.D. Hogan, G.P. Jones, L.J. McCook, M. Mills, R.L. Pressey, & D.H. Williamson. 2009. Connectivity, biodiversity conservation, and the design of marine reserve networks for coral reefs. Coral Reefs 28: 339-351. [ Links ]
Aronson, R.B. & W.F. Precht. 2001. White-band disease and the changing face of Caribbean coral reefs. Hydrobiologia 460: 25-38. [ Links ]
Aronson, R.B., I.G.MacIntyre, W.F. Precht, T.J.T. Murdoch & C.M. Wapnick. 2002. The expanding scale of species turnover events on coral reefs in Belize. Ecol. Monogr. 72: 233-249. [ Links ]
Babcock, R.C. 1991. Comparative demography of three species of scleractinian corals using age- and sizedependent classifications. Ecol. Monogr. 61:225–244. [ Links ]
Babcock, R. & C. Mundy.1996. Coral recruitment: consequences of settlement choice for early growth and survivorship in two scleractinians. J. Exp. Mar. Biol. Ecol. 206: 179-201. [ Links ]
Bak, R.P.M. & E.H. Meesters. 1998. Coral population structure: the hidden information of colony sizefrequency distributions. Mar. Ecol. Prog. Ser. 162: 301-306. [ Links ]
Bak, R.P.M. & E.H. Meesters. 1999. Population structure as a response of coral communities to global change. Am. Zool. 39: 56-65. [ Links ]
Ballantine, D.L., R.S. Appeldoorn, P. Yoshioka, E. Weil, R. Armstrong, J.R. Garcia, E. Otero, F. Pagan, C. Sherman, E.A. Hernandez-Delgado, A. Bruckner & C. Lilyestrom. 2008. Biology and ecology of Puerto Rican Coral Reefs. p. 375-406. In B.M. Reigl & R.E. Dodge (eds.) Coral Reefs of the World 1: Coral Reefs of the USA. Springer, Berlin, Germany. [ Links ]
Bellwood, D.R., T.P. Hughes, C. Folke & M. Nyström. 2004. Confronting the coral reef crisis. Nature 429: 827-833. [ Links ]
Bruckner, A.W. 2003. Proceedings of the Caribbean Acropora workshop: potential application of the Endangered Species Act as a conservation strategy. NOAA Technical Memorandum NMFS-OPR-24, Silver Spring, Maryland. [ Links ]
Bruckner, A.W. & R.J. Bruckner. 2003. Condition of coral reefs off less developed coastlines of Curaçao (stony corals and algae). Atoll Res. Bull. 496: 370-393. [ Links ]
Bruckner, A.W. & R.J. Bruckner. 2006a. Consequences of yellow band disease (YBD) on Montastraea annularis (species complex) populations on remote reefs off Mona Island, Puerto Rico. Dis. Aquatic Org. 69: 67-73. [ Links ]
Bruckner, A.W. & R.J. Bruckner. 2006b.The recent decline of Montastraea annularis (complex) coral populations in western Curacao: a cause for concern? Rev. Biol. Trop. 54 (Suppl. 1): 45-58. [ Links ]
Bruckner, A.W. & R. Hill. 2009. Ten years of change to coral communities off Mona and Desecheo Islands, Puerto Rico from disease and bleaching. Dis. Aquatic Org. 87: 19-31. [ Links ]
Bruno, J.F., E.R. Selig. 2007. Regional Decline of Coral Cover in the Indo-Pacific: Timing, Extent, and Subregional Comparisons. PLoS ONE 2(8): e711. doi:10.1371/journal.pone.0000711. [ Links ]
Bruno, J.F., E.R. Selig, K.S. Casey, C.A. Page, B.L.Willis, C.D. Harvell, H. Sweatman & A.M. Melendy. 2007. Thermal stress and coral cover as drivers of coral disease outbreaks. PLoS Biol. 5(6): e124. [ Links ]
Carpenter, R.C. 1990. Mass mortality of Diadema antillarum. I. Long- term effects on sea urchin populationdynamics and coral reef algal communities. Mar. Biol. 104: 67-77. [ Links ]
Carpenter, R.C. & P.J. Edmunds. 2006. Local and regional scale recovery of Diadema promotes recruitment of scleractinian corals. Ecol. Lett. 9: 271-80. [ Links ]
Cowen, R.K., C.B. Paris & A. Srinivasan. 2006 Scaling of connectivity in marine populations Science 311: 522-527. [ Links ]
Edmunds, P.J. & R.C. Carpenter. 2001. Recovery of Diadema antillarum reduces macroalgal cover and increases the abundance of juvenile corals on a Caribbean reef. Proc. Nat. Acad. Sci. USA 98: 5067-5071. [ Links ]
Edmunds, P.J. & R. Elahi. 2007. The demographics of a 15-year decline in cover of the Caribbean reef coral Montastraea annularis. Ecol. Monogr. 77: 3–18. [ Links ]
Folke, C., C.S. Holling, & C. Perrings 1996. Biological diversity, ecosystems, and the human scale. Ecol. Appl. 4: 1018–1024. [ Links ]
Gardner, T.A., I.M. Cote, F.A. Gill, A. Grant & A.R. Watkinson. 2003. Long-term region-wide declines in Caribbean corals. Science 301: 958-960. [ Links ]
Green, A.L. & D.R. Bellwood. 2009. Monitoring functional groups of herbivorous reef fishes as indicators of coral reef resilience – A practical guide for coral reef managers in the Asia Pacific region. IUCN working group on Climate Change and Coral Reefs. IUCN, Gland, Switzerland. 70 pages. [ Links ]
Grimsditch, G.D. & R.V. Salm. 2006. Coral reef resilience and resistance to bleaching. IUCN, Gland, Switzerland. 52 p. [ Links ]
Hall, V.R. & T.P. Hughes 1996. Reproductive strategies of modular organisms: comparative studies of reef building corals. Ecology 77: 950-963. [ Links ]
Harrison, P.L.& C.C. Wallace. 1990. Reproduction, dispersal and recruitment of scleractinian corals, p. 133-207 In Z. Dubinsky (ed.) Ecosystems of the world: coral reefs. Elsevier, Amsterdam, Holland. [ Links ]
Holling, C.S. 1973. Resilience and stability of ecological systems. Ann. Rev. Ecol. System. 4:1-23. [ Links ]
Hughes, T.P. 1994. Catastrophes, phase shifts, and largescale degradation of a Caribbean coral reef. Science 265: 1547–1551. [ Links ]
Hughes. T.P. & J.B.C. Jackson. 1985. Population dynamics and life histories of foliaceous coral. Ecol. Monogr. 55: 141-166. [ Links ]
Hughes, T.P., A.H. Baird, D.R. Bellwood, M. Card, S.R. Connolly, C. Folke, R. Grosberg, O. Hoegh-Guldberg, J.B.C. Jackson & J. Kleypas. 2003. Climate change, human impacts, and the resilience of coral reefs. Science 301: 929-933. [ Links ]
Hughes, T.P., D.R. Bellwood, C. Folke, R.S. Steneck & J. Wilson. 2005. New paradigms for supporting the resilience of marine ecosystems. Trends Ecol. Evol. 20: 380-386. [ Links ]
Hughes, T.P., M.J. Rodrigues, D.R. Bellwood, D. Ceccarelli, O. Hoegh-Guldberg, L. McCook, N. Moltschaniwskyj, M.S. Pratchett, R.S. Steneck & B. Willis. 2007. Phase shifts, herbivory, and the resilience of coral ceefs to climate change. Current Biol. 17: 360-365. [ Links ]
Jennings, S. & M.J. Kaiser. 1998. The effects of fishing on marine ecosystems. Adv. Mar. Biol. 34: 201-352. [ Links ]
Jones, G.P., G.R. Almany, G.D. Russ, P.F. Sale, R.S. Steneck, M.J.H. van Oppen & B.L. Willis. 2009. Larval retention and connectivity among populations of corals and reef fishes: history, advances and challenges. Coral Reefs 28: 307-325., N. & J.B.C. . . Shifting baselines, local impacts, and global change on coral reefs. PLoS Biology 6: e54, 6. [ Links ]
Lang, J.L. 2003. Status of coral reefs in the western Altantic: results of initial surveys, Atlantic and Gulf Rapid Reef Assessment (AGGRA) program. Atoll Res. Bull. 496, 630 p. [ Links ]
Lessios, H.A. 1988. Mass mortality of Diadema antillarum in the Caribbean: What have we learned? Ann. Rev. Ecol. Syst. 19: 371-393. [ Links ]
McClanahan, T.R., M.J. Marnane, J.E. Cinner & W.E. Kiene. 2006. A comparison of marine protected areas and alternative approaches to coral-reef management. Curr. Biol. 16: 1408-1413. [ Links ]
McClanahan, T.R., N.A.J. Graham, M. A. MacNeil, N.A. Muthiga, J.E. Cinner, J. Henrich Bruggemann & S.K. Wilson. 2011. Critical thresholds and tangible targets for ecosystem-based management of coral reef fisheries. Proc. Natl. Acad. Sci. USA 108: 17230-17233. [ Links ]
Meesters, E.H., M. Noordeloos & R.P.M. Bak. 1994. Damage and regeneration: links to growth in the reefbuilding coral Montastrea annularis. Mar. Ecol. Prog. Ser. 112:119-128. [ Links ]
Muller, E., C. Rogers, A. Spitzack & R. van Woesik. 2008. Bleaching increases likelihood of disease on Acropora palmata (Lamarck) in Hawksnest Bay, St John, US Virgin Islands. Coral Reefs 27: 191–195. [ Links ]
Mumby, P.J. 2006. The impact of exploiting grazers (Scaridae) on the dynamics of Caribbean coral reefs. Ecol Appl. 16: 747-69. [ Links ]
Mumby, P.J., C.P. Dahlgren, A.R. Harborne, C.V. Kappel, F. Micheli, D.R. Brumbaugh, K.E. Holmes, J.M. Mendes, K. Broad, J.N. Sanchirico, K. Buch, S. Box, R.W. Stoffle & A.B. Gill, 2006. Fishing, trophic cascades, and the process of grazing on coral reefs. Science 311: 98-101. [ Links ]
Mumby, P.J. & A.R. Harborne. 2010. Marine reserves enhance the recovery of corals on Caribbean reefs. PloS One 5(1): e8657. doi:10.1371/journal. pone.0008657. [ Links ]
Norström, A.V., M. Nyström, J. Lokrantz & C. Folke. 2009. Alternative states on coral reefs: beyond coral– macroalgal phase shifts. Mar. Ecol. Prog. Ser. 376: 295-306. [ Links ]
Nyström, M. & C. Folke. 2001. Spatial resilience of coral reefs. Ecosystems. 4: 406-417. [ Links ]
Nyström, M., N.A.J. Graham, J., Lokrantz & A. Norström. 2008. Capturing the cornerstones of coral reef resilience: linking theory to practice. Coral Reefs 27: 795-809. [ Links ]
Obura, D.O. & G. Grimsdith (eds.). 2009. Resilience Assessment of coral reefs – Assessment protocol for coral reefs, focusing on coral bleaching and thermal stress. IUCN working group on Climate Change and Coral Reefs. IUCN, Gland, Switzerland. 70 p. [ Links ]
Raymundo, L.J., C.S. Couch, A.W. Bruckner, C.D. Harvell, T.M. Work, E.Weil, C.M. Woodley, E. Jordan- Dahlgren, B.L. Willis, G.S. Aeby & Y. Sato. 2008. Coral Disease Handbook: Guidelines for Assessment, Monitoring and Management. CRTR, Australia. 131 p. [ Links ]
Roberts, C.M. 1997. Connectivity and management of Caribbean coral reefs. Science 278: 1454-1457. [ Links ]
Roberts, C.M., J.A. Bohnsack, F. Gell, J.P. Hawkins, and R. Goodridge, 2001. Effects of marine reserves on adjacent fisheries. Science 294: 1920-1923. [ Links ]
Rogers, C.S., J. Miller, E. Muller, P. Edmunds, R.S. Nemeth, J.P. Beets, A.M. Friedlander, T.B. Smith, R. Boulon, C.F.G. Jeffrey, C. Menza, C. Caldow, N. Idrisi, B. Kojis, M.E. Monaco, A. Spitzack, E.H. Gladfelter, J.C. Ogden, Z. Hillis-Starr, I. Lundgren, W.B. Schill, I.B. Kuffner, L.L. Richardson, B.E. Devine & J.D. Voss. 2008. Ecology of coral reefs in the US Virgin Islands, p. 303–374. In B. Riegl and R. E. Dodge (eds.) Coral Reefs of the World. 1. Coral Reefs of the USA. Springer, Berlin, Germany. Soong, K. 1991. Sexual reproductive patterns of shallowwater reef corals in Panama. Bull. Mar. Sci. 49: 832-846. [ Links ]
Soong, K. & J.C. Lang. 1992. Reproductive integration in reef corals. Biol. Bull. 183: 418-431. [ Links ]
Szmant-Froelich, A.M. 1985. The effect of colony size on the reproductive ability of the Caribbean coral Montastrea annularis (Ellis and Solander). Proc. 5th Int. Coral Reef Congress, Tahiti 4: 295-300. [ Links ]
Szmant, A.M. 1991. Sexual reproduction by the Caribbean reef corals Montastrea annularis and M. cavernosa. Mar. Ecol. Prog. Ser. 74: 13–25. [ Links ]
Sutherland, K., J. Porter & C. Torres. 2004. Disease and immunity in Caribbean and Indo- Pacific zooxanthellate corals. Mar. Ecol. Prog. Ser. 266: 273-302. [ Links ]
Weil, E. 2004. Coral reef diseases in the wider Caribbean, p. 35-68. In E. Rosenberg & Y. Loya (eds.) Coral Health and Disease. Springer-Verlag, Berlin, Germany. [ Links ]
Worm, B., E.B. Barbier, N. Beaumont, J. E Duffy, C. Folke, B.S. Halpern, J.B.C. Jackson, H.K. Lotze, F. Micheli, S.R. Palumbi, E. Sala, K.A. Selkoe, J.J. Stachowicz & R. Watson. 2006. Impacts of biodiversity loss on ocean ecosystem services. Science 314: 787-790. [ Links ]
*Correspondencia: Andrew W. Bruckner: Khaled bin Sultan Living Oceans Foundation, 8181 Professional Place, Landover, MD 20785 USA; Bruckner@livingoceansfoundation.org
1 Khaled bin Sultan Living Oceans Foundation, 8181 Professional Place, Landover, MD 20785 USA; Bruckner@livingoceansfoundation.org
Received 18-VII-2011. Corrected 12-XII-2011. Accepted 20-XII-2011.